Effects of Plants and Microorganism in Constructed Wetlands for Wastewater Treatment

www.elsevier.com/locate/biotechadv

Biotechnology Advances 22 (2003) 93–117

Effects of plants and microorganisms in constructed

wetlands for wastewater treatment

U. Stottmeister *, A. Wießner, P. Kuschk, U. Kappelmeyer,M. Kastner, O. Bederski, R.A. Muller, H. Moormann

UFZ Centre for Environmental Research, Leipzig-Halle, Germany

Abstract

Constructed wetlands are a natural alternative to technical methods of wastewater treatment.

However, our understanding of the complexprocesses causedby the plants,microorganisms, soilmatrix

and substances in the wastewater, and how they all interact with each other, is still rather incomplete.

In this article, a closer look will be taken at the mechanisms of both plants in constructed wetlands

and the microorganisms in the root zone which come into play when they remove contaminants from

wastewater. The supply of oxygen plays a crucial role in the activity and type of metabolism

performed by microorganisms in the root zone. Plants’ involvement in the input of oxygen into the

root zone, in the uptake of nutrients and in the direct degradation of pollutants as well as the role of

microorganisms are all examined in more detail.

The ways in which these processes act to treat wastewater are dealt with in the following order:

. Technological aspects;

. The effect of root growth on the soil matrix;

. Gas transport in helophytes and the release of oxygen into the rhizosphere;

. The uptake of inorganic compounds by plants;

. The uptake of organic pollutants by plants and their metabolism;

. The release of carbon compounds by plants;

. Factors affecting the elimination of pathogenic germs.

D 2003 Elsevier Inc. All rights reserved.

Keywords: Wetlands; Wastewater treatment; Microorganisms

0734-9750/$ - see front matter D 2003 Elsevier Inc. All rights reserved.

doi:10.1016/j.biotechadv.2003.08.010

* Corresponding author. Department of Environmental Biotechnology, UFZ Centre for Environmental

Research, Permoserstr. 15, D-04318 Leipzig, Germany. Tel.: +49-341-235-2220; fax: +49-341-235-2492.

E-mail address: [email protected] (U. Stottmeister).

U. Stottmeister et al. / Biotechnology Advances 22 (2003) 93–11794

1. Introduction

Treating wastewater in seminatural plant systems is a technique which can in principle

be applied in natural wetlands such as marshes, moors and wet fields, in artificial ponds

and lagoons, and in specially constructed wetlands. Constructed wetlands come in a

number of different basic designs featuring different flow characteristics (Kadlec, 1987;

Wissing, 1995).

The active reaction zone of constructed wetlands is the root zone (or rhizosphere). This

is where physicochemical and biological processes take place that are induced by the

interaction of plants, microorganisms, the soil and pollutants (Fig. 1).

Originally coined in 1903 by Hiltner and Stormer (1903), the term rhizosphere can be

subdivided into the endorhizosphere (the root interior) and the ectorhizosphere (the root’s

surroundings). The zone in which these two areas meet is known as the rhizoplane (Elliott

et al., 1984). This is where the most intensive interaction between the plant and

microorganisms is to be expected.

If wastewater is to be treated as efficiently as possible, detailed knowledge—such as the

effectiveness of various plant species, the colonization characteristics of certain groups of

microorganisms, and how biogenic compounds and particular contaminants (wastewater

components) interact with the filter bed material—is essential when designing constructed

wetlands. Research into constructed wetlands has chiefly dealt with technological design

issues, with the active reaction zone of the rhizosphere largely being treated as a ‘black box’

Fig. 1. Possible interactions in the root zone of wetlands for wastewater treatment.

Table 1

Selection of plant species used in constructed wetlands

Scientific name English name

Phragmites australis (Cav.) Trin. ex Steud. common reed

Juncus spp. rushes

Scirpus spp. bulrushes

Typha angustifolia L. narrow-leaved cattail

Typha latifolia L. broad-leaved cattail

Iris pseudacorus L. yellow flag

Acorus calamus L. sweet flag

Glyceria maxima (Hartm.) Holmb. reed grass

Carex spp. sedges

U. Stottmeister et al. / Biotechnology Advances 22 (2003) 93–117 95

where the only issues of concern were the inlet and outlet loads. This is almost solely

accounted for by the lack of suitable testing systems and study methods. However, small-

scale process modelling experiments are currently being developed (Kappelmeyer et al.,

2002).

According to practical experience and corresponding experiments, species of helo-

phytes (marsh plants) work best of all in seminatural wastewater treatment systems. This is

because helophytes possess specific characteristics of growth physiology that guarantee

their survival even under extreme rhizosphere conditions. The extreme conditions in the

rhizosphere in wetlands used to treat wastewater can be summed up as follow:

� Highly reduced milieu (Eh up to <� 200 mV, especially in horizontal subsurface flow

systems) prompting the formation of H2S and CH4;� Acidic or alkaline pH values in certain wastewaters;� Toxic wastewater components such as phenols, tensides, biocides, heavy metals, etc.;� Salinity.

Although all the plant species listed in Table 1 are suitable, reeds along with types of

rushes and cattails are the ones most frequently used. Recently, the suitability of fast-

growing trees such as willows has also been examined (Greenway and Bolton, 1996).

In order to learn more about the complexities of what happens when organic pollutants

are degraded in the root zone, we need to know more about the plants’ physiological

peculiarities and the microorganisms active in their rhizospheres.

2. Technological aspects

The knowledge accumulated over time about ways in which contaminants can be

removed by the simple natural combination of water, plants and soils has led to the

deliberate application of such systems in nature and ultimately to the creation of artificial

systems with various states of naturalness. Wissing (1995) divides the systems into the

following three main groups (see also Fig. 2):

� Aquaculture systems. Installations without active soil filters, such as ponds and ditches

with intensive growth of submerged aquatic and/or free-floating plants (Xu et al., 1992);

Fig. 2. Pond/wetland systems for wastewater treatment (A, pond with free-floating plants; B, horizontal surface

flow wetland or pond with emergent water plants; C, horizontal subsurface flow wetland; D, vertical flow

wetland).


� Hydrobotanical systems. Installations with a few active soil filters where removal is

mainly effected by aquatic plants, helophytes and microorganisms—ponds and ditches

with intensive growth of mainly helophytes;� Soil systems (see Fig. 2 for more details).

The basic types of soil-based constructed wetlands are:

� Horizontal surface flow systems (with the wastewater level above the soil surface);� Horizontal subsurface flow systems (with the wastewater level below the soil surface);� Vertical flow systems with upstream or downstream characteristics and continuous or

intermittent loading.

There are numerous different technological variants in terms of design, peripheral

equipment, etc. (Cooper, 1998). Usually, they are mainly distinguishable by the grain sizes

of the soil bed. In each case, the most suitable system can be adapted to specific waste

problems and local conditions. In addition, combination with other common methods of

waste pre- or posttreatment increases the possibilities available.

Mainly domestic wastewater, agricultural wastewater and mine drainage water are

treated in constructed wetlands (Mandi et al., 1998; Gearheart, 1992; Knight et al., 2000).

Increasing attention is now also being paid to using constructed wetlands to treat leachate,

contaminated groundwater and industrial effluents. The growing usage of such systems


has prompted intensive technological research and development in recent years. As far as

mainly domestic wastewater treatment is concerned, investigations have focused on the

following aspects:

� soil hydraulics (the influence of soil matter on hydraulic permeability), (Sanford et al.,

1995);� flow characteristics (horizontal or vertical flow, continuous or discontinuous water

load) (Netter, 1992; Stairs and Moore, 1994; Chazarenc et al., 2002);� external oxygen supply (by waterfalls, overflows and aeration installations, often in

connection with defined soil formations and discontinuous water loading or external air

input) (Green et al., 1998);� minimizing the area needed and maximizing the input load of contaminants; varying

the hydraulic retention time (Platzer and Netter, 1994; Platzer, 1999);� construction configuration, coupling different systems (Vymazal and Masa, 2002);� the plants used (empirical exploitation of different plants, mono vs. mixed culture;

influence of root growth) (Breen and Chick, 1995).

The precise technology chosen has an important influence on the contaminants’

biological degradation pathways and removal mechanisms. Whereas anaerobic processes

predominate in subsurface flow systems (apart from in the proximity of the helophyte

roots), aerobic processes usually prevail in surface flow systems. The hydraulic retention

time, including the length of time the water is in contact with the plant roots, affects the

extent to which the plant plays a significant role in the removal or breakdown of

pollutants. Whereas plants significantly affect the removal of pollutants in horizontal

subsurface systems with long hydraulic retention times used to clean municipal waste-

water, their role is minor in pollutant removal in periodically loaded vertical filters, which

usually have short hydraulic retention times (Wissing, 1995).

Following many years’ experience of working with constructed wetlands in countries

such as Germany and USA, operators have compiled manuals listing technical design

criteria and operating parameters (ATV, 1998; U.S. EPA, 1988; Water Pollution Control

Federation, 1990).

3. Root growth effect on the soil matrix

One important aspect of the complex processes taking place in the rhizosphere is the

interaction between roots/rhizomes and the soil matrix. The soil is the main supporting

material for plant growth and microbial films. Moreover, the soil matrix has a decisive

influence on the hydraulic processes.

Both chemical soil composition and physical parameters such as grain-size distribu-

tions, interstitial pore spaces, effective grain sizes, degrees of irregularity and the

coefficient of permeability are all important factors influencing the biotreatment system.

These physical parameters indicate certain hydraulic states of the soil and considerably

influence the flow of wastewater in constructed wetlands—and ultimately the removal of

contaminants. Root growth affects the physical (hydraulic) quality of soils (Kickuth, 1984;


Cooper and Boon, 1987; Wissing, 1995). On the one hand, roots and microbial biomass

clog up soil pores, but on the other hand, root growth and the microbial degradation of

dead roots cause the formation of new secondary soil pores.

As far as constructed wetlands are concerned, it seems that the main parameter

influencing the soil hydraulics is the grain-size distribution. Experience in Germany and

long-term studies of the hydraulics of constructed wetlands with different soil parameters

indicate that a mixture of sand and gravel produces the best results in terms of both

hydraulic conditions and the removal of contaminants (Wissing, 1995; Borner, 1990;

Netter, 1990).

For constructed wetlands with vertical flow, a relatively small range of effective grain

size d10 (the grain size below which 10 wt.% of the soil consists of, and which is the

leading characteristic of soils) from 0.06 to 0.1 mm was evaluated, while that for

constructed wetlands with horizontal flow was found to be higher at 0.1 mm (because

of the higher susceptibility to obstruction) (Wissing, 1995). The grain-size distribution of

>0.06 mm (up to 10 mm) with different distribution characteristics apparently enables

effective coefficients of permeability in the range of >10� 5 m/s (Wissing, 1995; Bahlo and

Wach, 1993), enough immobilization surface area for biofilm growth, positive impacts of

root growth on the hydraulic conductivity of the soil, and hence, all in all the good removal

of contaminants.

Investigations were conducted on a whole series of constructed wetlands which were

exclusively installed using soil material with Kf < 10� 8 m/s in order to maximize the area

for biofilm growth and the adsorption of wastewater chemicals (Morell, 1990; Bornert,

1990; Netter, 1990; Kretzschmar, 1990). The systems suffered hydraulic problems mainly

because of short-circuit flow on the wetland surface, and the predicted improvements in

the hydraulic conditions by root growth in the course of time were not observed. Despite

small increases in the Kf values (up to about 10� 7 m/s), the hydraulic conditions did not

change sufficiently. The main root growth of Phragmites australis was only recorded in

the top soil zone down to a depth of 20–30 cm (Bornert, 1990).

4. Gas transport in helophytes and oxygen release into the rhizosphere

In intermittently charged vertical filters, oxygen mainly enters the soil filter by virtue of

the suction effect of the water as it flows downwards. By contrast, in subsurface horizontal

flow systems oxygen is chiefly input by the marsh plants (helophytes).

How higher plants react to a lack of oxygen in the rhizosphere varies. Whereas typical

land plants which are adapted to dry locations cannot survive for long under such

conditions, plants which are adapted to waterlogged areas such as marshes, moors,

swamps and riverbanks have the anatomical and physiological attributes necessary for

their long-term survival (Vartapetian and Jackson, 1997).

The degree of adaptation is specific to individual species. Their survivability varies

over an extremely broad tolerance band from a few hours to several months, as

demonstrated by experiments in anaerobic incubators (Crawford and Braendle, 1996).

The fact that plants adapted to anoxic rhizosphere conditions can survive is because of

their ability to supply their root system with oxygen from the atmosphere. Gas transport


from the sections of the plant above the ground through the rhizome into the fine roots is

effected by specific areas of tissue formed in the plant known as the aerenchyma.

Depending on the degree of adaptation, these gas chambers can account for as much as

60% of the total tissue volume (Grosse and Schroder, 1986). The gas chambers in the

rhizome area are protected by node-like segmentation and diaphragms which are gas

permeable but nevertheless provide a secure barrier which prevents liquids from pene-

trating (Soukup et al., 2000). The various possibilities involved in the genesis of

aerenchyma structures by cell lysis or cell formation as well as their anatomical

peculiarities have been (and remain) the subject of thorough anatomical and physiological

studies (Jackson and Armstrong, 1999; Drew, 1997; Allen, 1997; Armstrong et al., 1994).

Interest is focused on issues such as how the interaction of changing environmental

conditions in the rhizosphere and biochemical processes causes anatomical changes in the

plant (Jackson and Armstrong, 1999).

The flow of gas through the plants is driven by diffusion processes and/or intensive

convective flows inducing high and low pressure (Allen, 1997; Armstrong et al., 1991;

Jackson and Armstrong, 1999; Grosse and Frick, 1999; Grosse, 1989). The types and

combination of the mechanisms involved are specific to each plant. For example, very

intensive convective gas transport has been observed in Typha latifolia (cattail) and P.

australis (reed) (Bendix et al., 1994; Armstrong and Armstrong, 1991). This convection is

caused by the formation of low pressure in oxygen-consuming sections of the plant and the

formation of higher pressure in the plant’s leaves (Allen, 1997). The formation of low

pressure is mainly based on the different solubilities of the oxygen used for restoration and

the carbon dioxide formed in this process. The formation of higher pressure in the leaves

causes air to flow throughout the entire body of the plant, with transport rates of up to 10

ml air per minute being measured (Grosse and Schroder, 1986; Schroder, 1986). One of

the main processes causing higher pressure is thermoosmosis (Grosse and Schroder, 1986;

Allen, 1997).

Owing to the differences in temperature between the cold phylloplane and the warmer

leaf interior, thermoosmosis causes air molecules to enter the young leaves through pores

(which are smaller than those in older leaves). The warmer interior of the leaf causes the

gas to expand owing to Brownian movement, limiting the possibility of returning through

the leaf pores. The overpressure building up inside the leaves is compensated for in the gas

transport tissue (the aerenchyma) inside the plant. As a result, the gas molecules are

transported through the plant right down to the deepest roots. The pressure compensation

of the plant system is finally achieved by gas being released through the roots and through

older leaves with larger pores.

The processes involved in the pressure-induced flow of gas in plants have been studied

since the mid-19th century (Grosse et al., 1996). This interest in the gas balance of higher

plants has been increasingly revived since the 1980s, not least in connection with the

growing interest in the biotechnological usage of plants adapted to flooded conditions, for

example, in order to clean wastewater (Grosse et al., 1996).

The introduction of atmospheric air into the plant’s interior means that under anoxic

conditions a sufficient amount of oxygen is available in the rhizome and root zones,

which can be used for respiration. However, the oxygen transported in the airflow is also

vital to the plant’s survival in another respect. Oxygen is released into the rhizosphere


and parts of the root system, mainly around the root tips and on young laterals

(Armstrong et al., 1990; Flessa, 1991). The release of oxygen causes the formation of

an oxidative protective film directly on the root surface. This film protects the sensitive

root areas from being damaged by toxic components in the anoxic, usually extremely

reduced rhizosphere (Armstrong et al., 1994; Vartapetian and Jackson, 1997). This

protective film has a thickness of between 1 and 4 mm depending on the way in which

incoming oxygen-consuming wastewater flows against the roots, and it contains redox

gradients ranging from about � 250 mV as frequently measured in reduced rhizospheres

to about + 500 mV directly on the root surface (Flessa, 1991). Oxygen is continuously

released from the internal root zones, counterbalancing chemical and biological oxygen

consumption.

This constant release of oxygen in the rhizosphere is of particular interest in connection

with the exploitation of the rhizosphere to treat wastewater. The oxygen flow rates which

have been measured, e.g., 126 Amol O2/h g root dry mass for Juncus ingens (giant rush)

(Sorrell and Armstrong, 1994) and 120–200 Amol O2/h g root dry mass for T. latifolia

(cattail) (Jespersen et al., 1998) are of biotechnological relevance. Model calculations for

P. australis (reed) resulted in area-specific oxygen input rates of 5–12 g O2/m2 patch area

per day (Armstrong et al., 1990). If the oxidation potential of certain helophytes is to be

put to optimal biotechnological use, this process must be quantified taking into account

the various factors of influence. These factors include rhizosphere-specific parameters

such as the redox state, pH, oxygen concentration, chemical characteristics and

temperature, plant-specific parameters such as mass and the species and stage of

development of plants, as well as phylloplane-specific factors such as temperature and

light intensity.

Studies have revealed that the redox state of the rhizosphere has a significant effect on

the intensity of oxygen release through the roots of various helophytes (Sorrell and

Armstrong, 1994; Sorrell, 1999; Kludze and Delaune, 1996; Wießner et al., 2002a). For

example, it was found that the release intensities clearly depend on the redox state of the

medium in the hydroponic vessel for T. latifolia and Juncus effusus plantlets under reduced

conditions, as shown in Fig. 3.

The oxygen release rates were highest at � 250 mV<Eh <� 150 mV. For extremely

reduced rhizospheric conditions (Eh <� 250 mV) and also moderately reduced rhizo-

spheric conditions (Eh>� 150 mV), the release intensities were found to be lower. The

potential of T. latifolia to release oxygen was much higher than that of J. effusus. Under

initially oxygen-free conditions but already in the positive redox range, all the plants of the

two species continued to release oxygen up to highly oxidized states.

Another interesting result of laboratory experiments is the correlation between root and

shoot size and oxygen release into the rhizosphere, as shown in Figs. 4 and 5. The relative

independence of the oxygen release rate from the root size and the significant influence of

the shoot size are evident.

For an explanation of this finding, it is well known that only small parts of the whole

root system (root tips and laterals) are permeable enough to transfer gases.

Otherwise, aspects of the aboveground biomass such as leaf areas and stomatal

conductance appear to be very important plant-specific parameters, including the plants’

capacities to release oxygen into their rhizosphere. The correlation between oxygen release

Fig. 3. Maximum oxygen release rates of individual experiments with several plantlets of T. latifolia, P. australis

and J. effusus depending on the corresponding redox potential (adapted from Wießner et al., 2002a, with

permission).


rates and shoot size can be described mathematically by nonlinear equations for both

species (see Fig. 6) (Wießner et al., 2002a).

Illumination was found to influence the intensity of oxygen release for T. latifolia

decisively but less so for J. effusus (Fig. 7). Obvious processes of aboveground gas

exchange and/or gas transport inside the plants (photosynthesis, thermoosmosis, diffusion,

Fig. 4. Oxygen release rates of J. effusus correlated with root and aboveground biomasses. ORR: oxygen release

rate; ADW: aboveground dry weight; RDW: root dry weight. ORR values are the mean of three experiments with

each plant (range of relative deviations 4–17%) (adapted from Wießner et al., 2002b, with permission).

Fig. 5. Oxygen release rates of J. effusus correlated with root and aboveground biomasses. ORR: oxygen release

rate; ADW: aboveground dry weight; RDW: root dry weight. ORR values are the mean of three experiments with

each plant (range of relative deviations 4–17%) (adapted from Wießner et al., 2002b, with permission).


pressure-induced gas flow) influenced by illumination affect the species-specific oxygen

release into the rhizosphere. Besides the redox state of the rhizosphere, the size and

physiological functionalities of the shoots were generally found to be of decisive

importance for supplying the rhizosphere with oxygen from the plants.

The aerenchyma tissue also plays a role in the emission of methane into the atmosphere

through emergent wetland plants (Thomas et al., 1996).

Mean rates of methane emission through helophyte plants in wetlands were estimated at

940 mg CH4/m2 day for a cattail wetland (Yavitt and Knapp, 1995), and up to 826 mg for a

rice field. Hence, in this rice, field over 95% of the methane emitted flowed through the

Fig. 6. Oxygen release rates of T. latifolia and J. effusus correlated to aboveground dry weight (adapted from

Wießner et al., 2002b, with permission).

Fig. 7. Oxygen release rates of T. latifolia and J. effusus correlated to illumination (adapted from Wießner et al.,

2002b, with permission).


rice plant (Banker et al., 1995). Thomas et al. (1996) summarized and cited other papers in

which helophytes are responsible for 50–90% of the total methane flux from wetlands.

Tanner et al. (1997) estimated methane emissions from constructed wetlands used to treat

agricultural wastewater to account for around 2–4% of wastewater carbon loads in

vegetated wetlands and 7–8% of loads in unvegetated systems.

Other pollutants (phytovolatilization) such as trichloroethylene and 2,6-dimethylphenol

were only emitted in trace amounts and are not significant for the technological treatment

process (Baeder-Bederski-Anteda, 2002).

5. The uptake of inorganic compounds by plants

The main mechanisms of nutrient removal from wastewater in constructed wetlands are

microbial processes such as nitrification and denitrification as well as physicochemical

processes such as the fixation of phosphate by iron and aluminum in the soil filter.

Moreover, plants are able to tolerate high concentrations of nutrients and heavy metals,

and in some cases even to accumulate them in their tissues.

The amount of phosphorous which can be taken up in the surface biomass of

Schoenoplectus lacustris (Sch. lacustris) is about 6.7 g m2 a� 1 (Seidel, 1966). The mean

phosphorous content in the dry biomass of a large number (41) of helophytes was found

by McJannet et al. (1995) to be around 0.15–1.05%. Consequently, less than 5% of the

phosphorus load in municipal wastewater is taken up by the plants. Seen from this angle,

the effect of harvesting the plant biomass is insignificant (Kim and Geary, 2001).

The uptake of nitrogen into the plant biomass is also of minor importance from a technical

viewpoint since harvesting the aboveground biomass would remove only 5–10% of the

nitrogen (Thable, 1984). Tanner (1996) estimated the nitrogen concentrations in helophytes

in the aboveground biomass to be between 15 and 32 mg N g� 1 dry mass. Owing to these

relatively low levels of nutrients, plant biomass is usually not harvested in Europe.


Wetlands—including constructed wetlands—are already being used to remove metals

from industrial effluents and mine drainage. Removal is chiefly based on the following

mechanisms:

� The oxidation of metals such as by iron-forming low-solubility precipitates in

oxygenated zones on the rhizome surface of helophytes in submerged systems (Wang

and Peverly, 1996) or in the free water zone of surface flow systems;� The coprecipitation of some elements such as arsenic with iron (ElbazPoulichet et al.,

2000);� Microbial sulfate reduction resulting in metal sulfide precipitates (Reynolds et al.,

1997);� The ion-exchanging capacity of the mineral and humic fractions of soil (Sobolewski,

1996);� Accumulation into plant matter (Dushenko et al., 1995).

From a technological viewpoint, the accumulation of heavy metals by plants is usually

insignificant when industrial effluent and mine drainage are being treated. This is because

the amount that can be accumulated is only a fraction of the total load of heavy metals in

wastewater. Nevertheless, a number of terrestrial plants are known which can accumulate

relatively high amounts of heavy metals in their biomass. Such plants are called ‘hyper-

accumulators.’ By definition, their dry biomass contains >0.1–1% metal (Baker, 1999).

Thio-reactive metals are sequestered in cysteine-rich peptides such as metallothioneins and

phytochelatins (Meagher, 2000). Elements such as arsenic, selenium and chromium are

subject to other mechanisms. For example, chromium(VI) is detoxified by reduction to

chromium(III) (Lytle et al., 1998).

At present, intensive research is being carried out to select hyperaccumulators which

are tolerant of heavy metals. In addition to natural breeding selection, new transgenic

plants are being developed (Macek et al., 2001). The aim is to develop inexpensive

techniques for ‘rhizofiltration’ (the removal of heavy metals and radionuclides etc. from

flowing wastewater) and ‘phytoextraction’ from soils.

6. Plant uptake and the metabolism of organic pollutants

One of the pioneers of using helophytes to treat wastewater was Seidel (1968). She

was the first person to study the removal of various phenols by helophytes in

hydroponic vessels as well as the tolerance of plants to phenols. Because these studies

were carried out as batch experiments under nonsterile conditions, the uptake is likely to

have been caused by both microorganisms and the plants themselves. In addition, no

constant test concentrations were established during her investigation of tolerance to

contaminants.

Infusion experiments with sterile plant tissue of Scirpus lacustris (Sc. lacustris) showed

an uptake of 0.08 mg phenol g� 1 fresh mass day� 1 (Kickuth, 1970). The main metabolite

identified was picolinic acid. It was therefore surmised that phenol degradation chiefly

takes place via catechol and further meta-ring cleavage. In the case of Lemna gibba,


phenyl-beta-D-glucopyranoside was identified as a metabolite of phenol degradation

(Barber et al., 1995).

Water hyacinth (Eichhornia crassipes) has often been studied in this regard. Wolverton

and McKown (1976) estimated the uptake of phenol to be about 36 mg/g dry substance

within 72 h. O’Keeffe et al. (1987a) studied the removal of various substituted phenols. The

uptake rate of the isomers decreased in the following sequence: paraHmeta > ortho cresol.

Toxicity increased with the rate of uptake. The acutely toxic phenol concentration was about

400 mg/l (O’Keeffe et al., 1987b), with catechol being identified as an early metabolite.

Important factors which influence the uptake of xenobiotics (organic pollutants) by the

plants include the compounds’ physicochemical characteristics such as the octanol–water

partition coefficient (log KOW), acidity constant (pKa), concentration, etc. (Wenzel et al.,

1999). Generally speaking, compounds with a log KOW between 0.5 and 3 are taken up best

(Trapp and Karlson, 2001).

Sandermann (1992) divides the metabolism of xenobiotics in plants into three phases:

� Transformation;� Conjugation;� Compartmentation.

The following enzymes are involved:

� Cytochrome P450;� Glutathione transferase;� Carboxylesterase;� O- and N-glucosyl transferase;� O- and N-malonyl transferase;

There are three possibilities for the final stage of detoxification:

� Export into the cell vacuole;� Export into the extracellular space;� Integration into lignin or other components of the cell membrane.

Despite the ability of plants to detoxify xenobiotics as described above, compared to

microorganisms they only play a secondary role in the direct degradation of organic

chemicals in wastewater treatment systems.

7. The release of carbon compounds from plants

The current knowledge about the input of carbon from plants into their rhizosphere

comes mainly from agricultural research.

The entire process of carbon input is known as rhizodeposition. Rhizodeposition

products (exsudates, mucigels, dead cell material, etc.) cause various biological processes

to take place in the rhizosphere. The quantity of organic carbon compounds released has

been estimated at 10–40% of the net photosynthetic production of agricultural crops


(Helal and Sauerbeck, 1989). The chemical composition of the exudates is very diverse.

Compounds which occur in plant tissues are usually also released through the roots. For

example, among the substances which have been identified in root exudates are sugars and

vitamins such as thiamine, riboflavin and pyridoxine etc., organic acids such as malate,

citrate, amino acids, benzoic acids, and phenol and other organic compounds (Miersch et

al., 1989). The range of substances varies from one species or even subspecies to the next.

It is assumed that the rhizodeposition products perform the following functions in the

rhizosphere:

� Mobilizing nutrients. Nutrient limitation can cause organic acids or other compounds to

be excreted. This may for example increase the solubility of iron and phosphate, thus

improving the plant’s nutrient supply (Hoffland et al., 1992).� Allelopathic effects. Some species of plants excrete special compounds into the

rhizosphere which impede the growth of other plant species (Miersch et al., 1989).

This effect has been examined in detail for a number of agricultural crops. The

literature does not yet contain any clear indications of allelopathy among helophytes

(Gopal and Goel, 1993).� Rhizosphere effects. Organic compounds such as sugars and amino acids can be used

by microorganisms as substrates, and excreted vitamins stimulate microbial growth.

Helal and Sauerbeck (1989) report that the majority of organic compounds excreted by

maize (80%) are mineralized by the microorganisms in the rhizosphere to form CO2,

increasing the microbial biomass in the rhizosphere. Furthermore, it has been shown

that organic compounds released by plants and plant residues influence the microbial

degradation of xenobiotics (Horswell et al., 1997; Donnelly et al., 1994; Moormann et

al., 2002). This issue is examined in more detail in Section 9.

Current knowledge of the composition of root exudates of helophytes is very limited, and

so far almost nothing is known about adult plants in this regard. Kaitzis (1970) investigated

rhizome extracts of Sc. lacustris and found various benzene derivatives including hydroxyl,

methoxyl, aldehyde and carboxyl groups. These extracted compounds displayed bactericidal

effects, indicating that they are responsible for the ‘negative’ rhizosphere effect (for more

details, see Section 10).

Owing to the relatively low amount of carbon released by plants in comparison to the

water flow, it can be assumed that rhizodeposition is only significant in constructed

wetlands if the carbon load in the wastewater is extremely low, as is for instance the case

with mine drainage. Rhizodeposition products can be used for bacterial dissimilatory

sulfate reduction. The H2S arising combines with heavy metal ions to form poorly soluble

sulfides in the anaerobic areas of the rhizosphere.

8. Transpiration

In addition to being of ecological importance, the transpiration of plants also influences

their technological application for wastewater treatment.


In practice, it is usually evapotranspiration that is measured. Evapotranspiration is the

sum of physical evaporation from water surfaces and plant transpiration.

The evapotranspiration rate varies sharply since it depends on numerous factors

influencing the ecosystem’s prevailing microclimate, as listed by Kadlec and Knight

(1996). For example, the values for tropical rainforest are about 1.5–2 m a� 1, compared to

about 0.4–0.5 m a� 1 for cornfields and forests in central Europe and 1.3–1.6 m a� 1 for

marshlands containing helophytes (Larcher, 1994).

In constructed wetlands used to treat wastewater in central Europe, water loss because

of evapotranspiration is about 5–15 mm day� 1 in the summer, i.e., some 20–50% of the

inflow (Schutte and Fehr, 1992). This aspect must be taken into account in warm periods

and in arid zones in order to prevent the water becoming excessively saline. When

choosing which technology to use under such extreme conditions, systems with low

hydraulic retention times (especially vertical filter systems) are to be preferred.

9. The role of the microbial degradation/transformation of organic and inorganic

pollutants

In constructed wetlands, the main role in the transformation and mineralization of

nutrients and organic pollutants is played not by plants but by microorganisms. Depending

on the oxygen input by helophytes and the availability of other electron acceptors, the

contaminants in the wastewater are metabolized in various ways. In subsurface flow

systems, aerobic processes only predominate near roots and on the rhizoplane (the surface

of the roots). In the zones that are largely free of oxygen, anaerobic processes such as

denitrification, sulfate reduction and/or methanogenesis take place.

Nitrogen transformation in constructed wetlands has already been the subject of several

papers. The main removal mechanism is microbial nitrification–denitrification; in con-

trast, incorporation into the plant biomass is of only minor importance (see Section 5).

Whereas in intermittently loaded vertical filters nitrate is often enriched, in subsurface

horizontal flow systems the oxidized nitrogen is immediately reduced, preventing the

enrichment of nitrite and nitrate.

Concerning subsurface horizontal flow systems, the nitrification step (forming nitrite or

nitrate) from which reduction takes place has not yet been determined. Furthermore, it is

not yet known whether anoxic ammonia oxidation according to the equation

5NHþ4 þ 3NO�

3 ! 4N2 þ 9H2Oþ 2Hþ

first specified by Van de Graaf et al. (1990) plays a significant part in this system.

In constructed wetlands, especially subsurface horizontal flow systems, very little

attention has been paid to the sulfur metabolism. In the case of an industrial wastewater

loaded with SO42� and S2O3

2� (area-specific load of 1.1 g S/m2 day), Winter (1985)

showed that constructed wetlands can act as an important sink for sulfur. Two percent of

the load was retained in the soil: 31% as S0, 25% as organic S (mainly in humic matter),

15% as sulfate and 11% as sulfide. Both microbial and abiotic processes are responsible

for these transformation processes.


Heavy metals are usually removed from industrial wastewater and mine drainage in

constructed wetlands by a variety of methods including:

� The filtration and sedimentation of suspended particles;� Adsorption;� Uptake into the plant material (see Section 5);� Precipitation by biogeochemical (microbial) processes.

In the aqueous phase of surface flow wetlands to treat mine drainage, Fe(II) is oxidized

to Fe(III) by abiotic and microbial oxidation; other elements such as arsenic also

precipitated. Other heavy metals are immobilized in the mainly anoxic soil by microbial

dissimilatory sulfate reduction and the H2S formed.

As far as highly chlorinated organic compounds are concerned, it is always much

easier for the corresponding carbon atom to be attacked by nucleophilic rather than by

oxidative reactions. Therefore, highly chlorinated hydrocarbons have a much higher

chance of being dehalogenated than less chlorinated compounds. Adrian et al. (1998)

demonstrated the microbial reduction of trichlorobenzene to form monochlorobenzene

via dichlorobenzene.

The only realistic possibility of biologically degrading hexachlorobenzene perchlo-

roethylene or highly chlorinated biphenyls is reductive dehalogenation (Wischnak and

Muller, 2000). The low-chlorinated products can then undergo further biological

degradation under aerobic conditions. Owing to the different redox states, existing as

Fig. 8. Influence of rhizodeposition products from P. arundinacea on 4-chlorophenol degradation by a mixed

culture obtained from P. arundinacea roots; meanF S.E., number of replicates: three. —.—, 4-chlorophenol

(reference); —E—, 4-chlorophenol + rhizodeposition products (12 mg DOC/l); —z—, 4-chlorophenol + rhi-

zodeposition products (4 mg DOC/l) (adapted from Moormann et al., 2002; with permission).


a result of the oxygen-donating helophyte roots, a constructed wetland is a metabolically

multipotent ‘technical ecosystem.’ Hence, the abovementioned conditions ought to

enable the microbial degradation of both highly chlorinated and low-chlorinated

compounds in this ecosystem.

It is also conceivable that in zones of constructed wetlands with a low organic

load, root exudates and dead plant material could be involved in the microbial

cometabolic degradation of poorly degradable organic compounds (Moormann et al.,

2002).

Whereas rhizodeposition products obtained from helophytes have not been observed

to have any enhancing biodegradation effects on phenol or 2,6-dimethylphenol, a

stimulating effect was found in experiments with the more recalcitrant 4-chlorophenol.

For example, the degradation of 4-chlorophenol by a mixed bacterial culture obtained

from Phalaris arundinacea roots was enhanced by the rhizodeposition products (Fig. 8).

Subsequent experiments with characterized bacteria from the plant in pure culture

confirmed the effect of accelerating degradation (Acinetobacter baumannii and especially

Ralstonia sp.).

The function of rhizodeposition products as growth substrates for 4-chlorophenol

degradation was confirmed with Ralstonia eutropha (DSMZ strain 5536). For R. eutropha,

the cometabolism of 4-chlorophenol with phenol as a growth substrate has already been

described (Hill et al., 1996).

10. Factors affecting the elimination of pathogenic germs

It was back in the 1970s that Seidel (1971, 1972, 1973a,b) first drew attention to the

bactericidal effect of higher plants on pathogenic germs. She put this aspect to practical

use in the first constructed wetlands built in Germany. Her laboratory experiments revealed

that the effect varied depending on the species of plant used. For example, 10 plant species

managed to remove 99% of Escherichia coli in pot experiments within 48 h, while 13

species removed 85% and another 31 species only eliminated 15%. Among the helo-

phytes, Mentha aquatica, Alisma plantago and J. effusus proved to be especially efficient

(Seidel, 1971). Seidel’s findings were largely confirmed by Burger and Weise (1984) in

pot experiments using 1.2 l of sand and 5 l of nutrient solution. In pots containing Glyceria

maxima, S. lacustris, A. plantago-aquatica and M. aquatica, the number of bacteria

(colony-forming units) was reduced by 90% after a contact time of 7–11 h and by 99%

after 16–19 h. The effectiveness was highest in the first few days of the investigation

period. Compared to the control experiments, the reduction time was between a third and a

half shorter.

Unfortunately, the findings of the two groups (Seidel vs. Burger and Weise) cannot be

directly compared since the descriptions of the experimental method (the way in which the

experiments were carried out, the state of the plants, etc.) are too incomplete.

Vincent et al. (1994) studied the bactericidal effect of the in vitro (aseptically)

cultivated helophytes M. aquatica, P. australis and Sc. lacustris on E. coli. In the test

system, all three plant species inhibited the growth of E. coli, with the strongest effect

being exhibited by Sc. lacustris. Only in the case of M. aquatica did plant-free nutrient


solutions from which the plants had been removed after 14 days’ cultivation show a slight

antibacterial effect. The authors concluded that the bactericidal effect is an active process

that requires the direct presence of plants.

Despite these findings, it is difficult to explain the fact that the root exudates both

stimulate and inhibit bacterial growth. Therefore, other mechanisms and indirect effects by

plants (adsorption, aggregation and filtration) as well as the effect of protozoa have been

discussed (Kadlec and Knight, 1996).

The exact role of biolytic processes (such as the action of protozoa, bdellovibrios and

bacteriophages) in germ reduction in constructed wetlands is still largely unknown.

Gradl et al. (1994) carried out small-scale experiments into the removal of fecal coli and

total coli from the discharge of a mechanical–biological clarification plant with a

vertically charged soil filter system (2.5 m/day). The filter was planted with Acorus

calamus and P. australis. The filter planted with A. calamus achieved bathwater quality in

terms of fecal coli and total coli. Unfortunately, this paper does not allow any scientifically

sound conclusions. The effectiveness of filters planted with A. calamus and P. australis

was compared by using findings from various years; the state of planting was not

characterized, and no unplanted control was used by way of comparison.

Rivera et al. (1995) confirmed the effect of the increased elimination of bacteria (E.

coli) in the rhizosphere by Phragmites and Typha (35–91%) compared to unplanted

controls in microcosm investigations. However, no significant efficiency differences were

found between the two plant species (Phragmites and Typha). During the pilot-scale tests,

a lower rate of elimination was noted in the winter than in the summer, although no

significant differences were observed between the planted and unplanted variants.

In an Austrian study, the disinfection parameters were characterized in three vertically

charged constructed wetlands (Mitterer, 1995). In terms of colony-forming units, elimina-

tion was in the order of between three and four orders of magnitude. The rates of elimination

for fecal coliforms and enterococci were 3–4 and 2–3 orders of magnitude, respectively.

Kadlec and Knight (1996) listed the efficiency of the elimination of coliforms and

streptococci in various systems of constructed wetlands. As a rule, more than 90% of the

coliforms and more than 80% of the fecal streptococci were eliminated.

Hagendorf and Hahn (1994) studied the efficiency of a number of wetlands in

Germany. They observed the best results in systems with a mixture of sand and gravel

and vertical flow. Horizontal systems were by no means as efficient, although those with

fine to medium sandy soil afforded better germ reduction than those with pebbly soil.

However, systems with small-grained soils often resulted in hydraulic problems (clogging

leading to surface flow), which drastically reduced efficiency.

The findings of Thurston et al. (1996) regarding the comparison of a pond system with

a subsurface flow planted soil filter are very interesting, and are listed in Table 2.

As expected, the planted soil filter is more efficient at eliminating bacteria than the

Lemna pond. Surprisingly, however, the protozoa were better eliminated in the pond than

in the soil filter.

Rivera et al. (1995) also observed that the elimination of pathogens cannot be solely

explained by filtration effects. Whereas amoebae in a gravel filter were removed to a

degree of 95% (tropical climate) or 75% (subtropical climate), the efficiency of the soil

filter was only 15–17%.

Table 2

Comparison of the efficiency of a pond and a planted soil filter for wastewater disinfection (Thurston et al., 1996)

(elimination efficiency in %)

Lemna pond Planted soil filter

Giardia cysts 93 83

Cryptosporidium oocysts 91 67

Total coliforms 54 99

Fecal coliforms 59 98

Coliphages 35 94


As the above examples show, the efficiency of germ elimination in constructed

wetlands is subject to sharp fluctuation. In contrast to systems with low cleaning

performance, examples are known in which the epidemic disinfection standard for

agricultural irrigation and aquaculture contained in WHO guidelines of V 1000 fecal

coli/100 ml (WHO, 1989) are met (Green et al., 1997).

These positive examples demonstrate the potential of this cleaning technology, even if

the mechanisms of germ reduction are not fully understood.

The very complex mechanisms in these systems have so far only been studied to a

limited extent. According to Ottova et al. (1997), important factors of influence in

connection with germ reduction include the following:

� Physical: filtration, sedimentation, adsorption and aggregation;� Biological: consumed by protozoa, lytic bacteria, bacteriophages, natural death;� Chemical: oxidative damage, influence of toxins from other microorganisms and

plants;

11. Outlook

Constructed wetlands have been used to treat wastewater ever since the pioneering

work performed by Kathe Seidel in the 1960s.

Over the years, numerous examples have shown that this technology is suitable for

treating both municipal sewage and a broad range of industrial wastewater. It also gave

birth to the idea of using trees to remediate contaminated aquifers near the surface.

Whereas, originally, merely small wetlands were constructed which could only treat the

wastewater produced by a small population, in the meantime, larger treatment works have

also been built that are able to cope with a population equivalent of several thousands.

Therefore, the consequences of using this technology should be examined, such as the

technological limits of wetland size and the problems which are to be expected with large

constructed wetlands. In the past, for example, sewage farms (farmland irrigated with

wastewater) were used, which resulted in the accumulation of pollutants over a long

period. This problem can be minimized by the separate treatment of industrial and

municipal wastewater accompanied by careful monitoring. Another question that needs

to be answered is the extent to which the plant biomass in such large wetlands can

subsequently be put to viable economic use, for example, as a source of energy or as raw


material for the paper industry, etc. These are issues that will naturally vary from one

country to the next depending on the various socioeconomic and climatic conditions that

apply.

Given the pressing need for clean, disinfected water in the ‘Third World,’ this is an area

where the usage of constructed wetlands to solve water and wastewater problems could be

especially widespread. This is a specific field which needs further research and develop-

ment work.

In future, once constructed wetlands become better known, increasing attention is likely

to be paid to developing suitable combinations of different technologies. For example, one

energy-saving strategy would be the combination of anaerobic fermentation and post-

treatment in constructed wetlands.

Despite the experience which has been built up in years of practical application and

research, a number of fundamental aspects of exactly how constructed wetlands function

are not yet adequately understood. One reason for this is that, compared to other

technologies such as activated sludge, constructed wetlands depend on the interaction of

many more different components.

The basic aspects upon which more work is essential include:

� The microbial process of anoxic ammonium oxidation and possibilities of stimulating it

in constructed wetlands;� The behavior of toxicologically highly active trace substances such as persistent drug

residues in the complex system of the constructed wetland, which theoretically ought to

allow better cleaning effectiveness;� The behavior of persistent compounds in connection with how they can be removed

from wastewater in this complex system and how they can either be detoxified or made

safe by means of their long-term immobilization;� The mechanisms of wastewater disinfection, with particular attention to the role of

biolytic processes;� Although the effect of root growth on hydraulic conductivity (especially in small-

grained soil) has been much discussed, so far, only a few plant species have been

investigated in this regard;� The use of plants, which have been adapted to a specific wastewater problem, will

continue to remain an important topic of future research. Genetic engineering provides

a growing number of methods to breed plants, which can, for instance, better

accumulate heavy metals or break down persistent contaminants more effectively in

constructed wetlands. However, attention must always be paid to how these ‘new’ plant

species can stand up in the long term to the many competing influences such as wild

species in these complex technical ecosystems—just as crops in the field permanently

have to compete with weeds.� The interactions of various substance cycles (e.g. carbon, nitrogen, and sulfur), taking

into account above all the variability of redox states in the rhizosphere.

Achieving a better understanding of the complex interactions involved will enable the

basic scientific aspects to be optimally combined with the technical possibilities available,

thus enabling wetland technologies to be used on a broader scale.


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Effects of Plants and Microorganism in Constructed Wetlands for Wastewater Treatment

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